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Comments of the
European Council for Alkylphenols and Derivatives
and the
Alkylphenols & Ethoxylates Research Council
On the REACH Annex XV Report: Proposal for Identification of a Substance as a CMR
Cat. 1A or 1B, PBT or vPvB, or a Substance of an Equivalent Level of Concern:
4-Nonylphenol, branched and linear
Submitted October 17, 2012
Executive Summary
The European Council for Alkylphenols and Derivatives (CEPAD) and the Alkylphenols &
Ethoxylates Research Council (APERC) jointly submit these comments in objection to the
Annex XV Report on 4-nonylphenol(4-NP), which proposes classification of this compound as a
Substance of Very High Concern (SVHC) under Regulation (EC) 1907/2006 (REACH).
The criteria for SVHC are listed under Article 57as:
(a) substances meeting the criteria for classification as carcinogenic category 1 or 2 in
accordance with Directive 67/548/EEC;
(b) substances meeting the criteria for classification as mutagenic category 1 or 2 in
accordance with Directive 67/548/EEC;
(c) substances meeting the criteria for classification as toxic for reproduction category 1
or 2 in accordance with Directive 67/548/EEC;
(d) substances which are persistent, bioaccumulative and toxic in accordance with the
criteria set out in Annex XIII of REACH;
(e) substances which are very persistent and very bioaccumulative in accordance with the
criteria set out in Annex XIII of this Regulation; and
f) substances — such as those having endocrine disrupting properties or those having
persistent, bioaccumulative and toxic properties or very persistent and very
bioaccumulative properties, which do not fulfill the criteria of points (d) or (e) — for
which there is scientific evidence of probable serious effects to human health or the
environment which give rise to an equivalent level of concern to those of other
substances listed in points (a) to (e) and which are identified on a case-by-case basis in
accordance with the procedure set out in Article.” (emphasis added)
The Annex XV Report for 4-NP proposes these compounds as SVHC “because they are
substances with endocrine disrupting properties for which there is scientific evidence of probable
serious effects to the environment which give rise to an equivalent level of concern to those of
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other substances listed in points (a) to (e) of Article 57 of REACH.” The proposal is based
primarily on the following arguments cited in the summary of the Report (pg. 5):
1.) “There is “strong evidence from high quality studies of endocrine mediated
adverse effects in fish species”; and
2.) “Results for amphibians provide indications that effects in other taxa may be
endocrine mediated i.e. caused by an estrogen-like mode of action, too.”
These comments will address why these two arguments, and others provided in the Annex XV
Report for 4-NP, are scientifically inadequate to justify identifying this compound as SVHC.
In addition to scientific concerns, CEPAD and APERC have process concerns with the fact that
there are currently 54 substances nominated for identification as SVHC under public consultation
with the same deadline for comment on October 18, 2012. Managing such a large number of
dossiers within a short time frame will certainly be a challenge for both the ECHA Secretariat
and the Member State Committee, which raises questions about the how thoroughly public
comments will be considered.
Also, CEPAD and APERC object to the proposal to nominate 4-NP as SVHC because it raises a
fundamental policy issues that require further clarification under REACH before proceeding with
SVHC nominations under Article 57(f). There are currently no criteria for what constitutes
“probable serious effects to the environment” from endocrine modes of action, or any other
mode of action, established under REACH. In fact, the EU Commission still has ongoing
activities related to the development of a definition of “endocrine disruptors” and criteria for
their assessment under REACH. Therefore, it is premature to consider nominations for
compounds as SVHC on this basis in advance of these policy developments at the level of the
European Commission.
4-NP does not meet the criteria for SVHC under Article 57(a) through (e). 4-NP does not meet
the criteria for PBT or vPvB as determined by several governmental authorities. (Joint Meeting
of the Competent Authorities for the Implementation of Council Directive 67/548/EEC, 2001;
Joint Research Centre, Institute for Health and Consumer Protection, TC NES Subgroup on
identification of PBT and vPvB Substances). 4-NP also does not meet the criteria for probable
serious effects for SVHCs as defined under CMR Cat. 1 or 2; it also does not meet criteria for the
weakest category, CMR Cat. 4. (ECB, 2002) However, 4-NP is toxic to aquatic organisms. As
discussed later in these comments, the estrogenic mode of action is not the sole mode of action
for 4-NP, and the most sensitive aquatic ecotoxicity end points for 4-NP are not definitively
linked to an estrogenic mode of action. Furthermore, considering that the adverse effects of 4-NP
on aquatic organisms is already controlled through existing regulatory instruments such as the
Water Framework Directive, which has an Environmental Quality Standard (EQS) for 4-NP, and
the Integrated Pollution Prevention and Control (IPPC) Directive, there does not appear to be a
need to further prioritize a nomination for this compound as SVHC on the basis of concern for
the toxicity to the environment
In addition, there is a long-standing EU Market and Use Directive (M&U Directive) for
Nonylphenol and its Ethoxylates first issued under DIRECTIVE 2003/53/EC OF THE
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EUROPEAN PARLIAMENT AND OF THE COUNCIL of 18 June 2003 amending for the 26th
time Council Directive 76/769/EEC and now incorporated into the REACH regulation as (EC)
No 1907/2006, Annex XVIII. The M&U Directive for 4-NP/NPE was based on concern for the
aquatic toxicity of 4-NP and restricted the marketing and use of 4-NP and its ethoxylates in the
EU for major wide dispersive uses that result in a direct discharge to the aquatic environment.
In the absence of EU criteria, the Annex XV Report for 4-NP proposes that this compound meets
the threshold for “equivalent level of concern” under Article 57(f); however it does not provide
an adequate scientific basis to support a case that this compound rises to a level of concern that is
equivalent to a CMR Category 1, 2, PBT or vPvB compound. Article 57(f) of REACH states
“substances giving rise to an equivalent level of concern” should have “scientific evidence of
probable serious effects to human health or the environment”; therefore it seems that the
intention of the regulation is that SVHC should be reserved for substances for which there is an
adequate scientific basis to support a finding of effects that are both probable and serious.
As will be discussed further in these comments, 4-NP has only weak estrogenic activity that is
1,000 to 1,000,000 fold less potent than the potent estrogens17ß-estradiol (E2) and 17α-
ethynylestradiol (EE2), when calculated based on numerous reliable in vitro and in vivo studies.
Adverse apical effects caused by 4-NP are not "clearly endocrine mediated", but rather are
indicative of general toxicity (baseline narcosis) possibly coupled with very weak estrogenic
activity. The effects of 4-NP are clearly not comparable to other more potent steroidal
estrogens. Results of aquatic toxicity tests conducted on 4-NP and potent estrogens show that
effects on acute lethality, reproduction, growth, and development are very different for 4-NP and
EE2.
The following comments elaborate on these points and respond to the specific arguments
provided in the Annex XV Report for 4-NP to demonstrate that while 4-NP may have some weak
estrogenic activity, it does not demonstrate adverse environmental effects that give rise to an
equivalent level of concern as CMR, PBT or vPvB compounds.
1.0 The IPCS Definitions distinguish between “endocrine disruptor” and “potential
endocrine disruptor”; and the OECD Guidance Document on Standardised Test
Guidelines for Evaluating Chemicals for Endocrine Disruption provides testing
guidance, not policy guidance for regulatory prioritization or action related to
endocrine active compounds.
As noted above, CEPAD and APERC object to the proposal to nominate 4-NP as SVHC because
it raises a fundamental policy issues that require further clarification under REACH before
proceeding with SVHC nominations under Article 57(f). There are currently no criteria for what
constitutes “probable serious effects to the environment” from endocrine modes of action, or any
other mode of action, established under REACH. In fact, the EU Commission still has ongoing
activities related to the development of a definition of “endocrine disruptors” and criteria for
their assessment under REACH. These activities include discussion about an endocrine disruptor
hazard classification scheme and consideration of the potency, adversity of effects and
requirement to review the weight-of-evidence for endocrine disruptors.(Kortenkamp, 2011).
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In the absence of EU definitions and guidance on the “probable serious effects to the
environment” from endocrine modes of action, the Annex XV Report for 4-NP relies on the
IPCS definition and the OECD Guidance Document Guidance Document on Standardised Test
Guidelines for Evaluating Chemicals for Endocrine Disruption. (OECD, 2011) However, the
Report overlooks the following features of these sources on endocrine disruption.
1.1 The IPCS Global Assessment of Endocrine Disrupting Compounds distinguishes
between an “endocrine disruptor” and a “potential endocrine disruptor”.
The IPCS in their Global Assessment of Endocrine Disrupting Compounds (EDCs) stated that:
“An endocrine disruptor is an exogenous substance or mixture that alters function(s) of
the endocrine system and consequently causes adverse health effects in an intact
organism, or its progeny, or (sub)populations; and
A potential endocrine disruptor is an exogenous substance or mixture that possesses
properties that might be expected to lead to endocrine disruption in an intact organism,
or its progeny, or (sub)populations.” [emphasis added] (WHO, IPCS,2002).
Most noteworthy, for a chemical to be considered an endocrine disruptor, the endocrine mediated
effects must occur in a whole organism and they must be adverse. IPCS stresses that: “Endocrine
disruption is not considered a toxicological end point per se but a functional change that may
lead to adverse effects.” Thus, chemicals that show some endocrine modulation (e.g.,
estrogenic, androgenic, thyroidogenic) are not necessarily endocrine disruptors and should not be
considered to be such.
The U.S. Environmental Protection Agency makes this point in their two-tiered Endocrine
Disruption Screening Program (EDSP), where results from Tier 1 screening tests (which may
indicate potential endocrine activity) are not indications of definitive “endocrine disruption”.
Rather results from Tier 2 multigenerational toxicity tests with intact organisms determine
whether a substance may cause endocrine-mediated effects through or involving various
hormone systems. (US EPA, EDSP, 2012)
It should be noted that the WHO-IPCS report recognizes that not all endocrine active compounds
should be considered as endocrine disruptors. Certainly, the potency, efficacy, as well as
exposure concentrations, will affect the probability that a compound will actually cause serious
endocrine disruption. In addition, the influence of other co-occurring modes of action will
determine whether endocrine activity or disruption is the primary mode of action. The Annex
XV Report for 4-NP overlooks these aspects of the IPCS Global Assessment of Endocrine
Disrupting Compounds.
1.2 The OECD Guidance Document Guidance Document on Standardised Test
Guidelines for Evaluating Chemicals for Endocrine Disruption is intended as a tool
to assist regulatory authorities to interpret assays and explicitly states that it is not
intended to prejudge or constrain regulatory actions in its testing guidance.
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The objectives and scope of the OECD Guidance Document were defined such that the
document would be a tool to support regulatory authorities by helping to interpret assay results
and suggesting possible additional studies for reducing uncertainty. The Guidance Document
specifically states that “the guidance should not prejudge or constrain what regulatory actions
may be taken by a member country and should not suggest a testing strategy”. (OECD, 2011).
The Guidance Document also notes “the use of many of these tests for determination of toxicity
due to endocrine disruption (hazard and risk assessment) for mammals and wildlife is rather new,
and therefore the guidance given is considered to be subject to changes based on new evidence”.
The Guidance Document is intended to be a “living” document that will be updated as the
science in this area evolves and should not be relied upon as a regulatory standard or criteria.
In the OECD Guidance Document, an endocrine disrupter (ED) has been defined according to
WHO IPCS, 2003 as described above, and “it is acknowledged that many other definitions exist
(e.g. Weybridge Conference, 1996) but the WHO (2003) definition has been used as a working
definition for this document because it covers both human health and wildlife populations”. The
Guidance Document points out that this definition is widely used but not universally accepted.
The OECD Guidance Document “operationally defined the term ‘possible ED’ to mean “a
chemical that is able to alter the functioning of the endocrine system but for which information
about possible adverse consequences of that alteration in an intact organism is uncertain”.
This raises the need for the development of EU criteria for determining probable and serious
effects in aquatic species, regardless of mode of action. Endocrine mediated reproductive effects
that are “known” or “presumed” based on supporting data that show they occur in the absence of
other interfering toxicological mechanisms may rise to a similar level concern as Cat. 1 or 2
(based on DSD classification criteria) reproductive toxicant, particularly when effects occur at
exposures that are environmentally relevant. However, reproductive effects that occur at only
concentrations that are so high that other modes of toxicity are also occurring or that are not
environmentally relevant should not be considered to be of equivalent concern to CMR Cat. 1, 2
or PBT or vPvB compounds.
2.0 4-NP is very weakly estrogenic, operates by multiple modes of action, and does not
result in serious adverse endocrine-mediated effects comparable to the potent
natural estrogen, 17β-estradiol (E2) and synthetic (EE2) estrogens.
2.1 4-NP is considerably less potent than natural (E2) and synthetic (EE2)
estrogens based on in vitro estrogen activity screening tests and endpoints
from in vivo tests.
Valuable information about the relative potency of endocrine active compounds relative to the
known potent estrogens E2 and EE2 are provided by various in vitro screening tests and in
vivo tests with intact organisms. Endpoints from in vivo tests included gonadal histology,
kidney lesions, vitellogenin, sex ratio, hatching success, and swim-up. Relative potencies from
these studies can be approximated by concentrations of 4-NP and E2 or EE2 used as a positive
control that gave comparable results. Overall, the data show that 4-NP is approximately 103–
106-fold less potent than the endogenous estrogen E2 or the synthetic estrogen EE2, depending
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on the species and endpoint investigated (Jobling and Sumpter 1993; Lee and Lee 1996;
Islinger et al. 1999; Metcalfe et al. 2001; Dussault et al. 2005; Balch and Metcalfe 2006; Lin
and Janz 2006; Zha et al. 2007, 2008).
2.2 Examination of the estrogenic potency and aquatic toxicity of natural estrogen
(E2) and synthetic estrogen (EE2) provides an understanding of what constitutes
a highly potent estrogen.
The Annex XV Report on 4-NP emphasizes the estrogenic activity of 4-NP as the basis for
“equivalent concern” in its assessment of the compound’s aquatic toxicity. In light of this
concern, it is of use to first examine effects on aquatic organisms from known endocrine active
substances for which the mode of action is indisputable – natural estrogen (E2) and a synthetic
and therapeutic estrogen (EE2). Relevant studies are summarized in (Table 2) and more fully
presented with experimental details in the attached Table A.
Table 2. Summary of Aquatic Toxicity Studies for 17-ethynylestradiol (EE2) and 17-estradiol (E2)
Study Type NOEC Apical
Endpoints
NOEC Secondary
Endpoints
Comments References
Screening or
Short Term
Reproduction
Fathead minnow
Sheepshead
minnow
Zebrafish
Survival,
reproduction,
fertilization spawning
success, hatch
success, sex ratio:
≤0.005 to 0.20 g/L
VTG, GSI, kidney
lesions, gonadal
histopathology,
testis-ova:
0.0002 to 0.010 g/L
Spawning,
fertilization,
mating behavior
most affected.
Effects on gonadal
histopathology
severe
Miles-
Richardson
1999
Zilloux 2001
Van den Belt
2001
Coe 2010
Life Cycle
Studies
Fathead minnow
Zebrafish
F0 fertilized eggs to
adult exposure,
survival, growth,
reproduction:
0.0001 to 0.016 g/L
F1 sex ratio:
0.0002 g/L
Secondary sex
characteristics, testis-
ova, gonadal
histopathology, VTG:
0.001 to 0.004 g/L
Complete
feminization of
adult fish.
No male external
characteristics or
gonadal tissue.
Lange 2001
F0 fertilized eggs to
adult exposure,
survival, growth,
reproduction:
0.0003 to 0.010 g/L
F1 fertilization:
0.0003 g/L
Not reported Ratio of acute
LC50 to
fertilization NOEC:
1700 g/L/0.0003
g/L =
5.73 million.
Wenzel 2001
Whole Lake
Studies
Fathead minnow
Pearl dace
Whole lake exposed
for 3 years to 0.005
g/L EE2
Complete collapse of
fathead minnow
populations.
Decreased abundance
of pearl dace
populations.
VTG
Edema in ovaries
Inhibition of male
gonad development,
intersex tissues,
kidney lesions
Complete
feminization of
males, inability to
reproduce.
Kidd 2007
Palace 2006
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E2 and EE2 have been found to be of generally similar potencies to each other. EE2/E2 ratios
range from 0.19 to 1.9 depending on the screen used (Kidd et al., 2007). E2 and EE2 are
especially potent when compared to the weak estrogenic activity of 4-NP.
The natural and synthetic estrogens E2 and EE2, respectively, cause significant and severe
effects on freshwater and saltwater fish. As described in more detail below, a signature set of
aquatic effects from E2 and EE2 may include: complete feminization of male gonads;
induced failure to differentiate into males; reduced reproduction capabilities; severe
histological effects on gonadal tissues; and even collapse of whole populations in a lake
continuously dosed with EE2 for three years.
A series of short term reproduction assays showed that survival, reproduction, fertilization,
spawning and hatch success, and sex ratio can be significantly altered at estrogen
concentrations down to 0.005 µg/L. Secondary endpoints such as gonadal weights and
histopathology, testis-ova, and kidney lesions were affected at even lower concentrations,
down to 0.0002 µg/L.
Life cycle studies with fathead minnow and zebrafish that included exposure to EE2 over the
full life cycle showed similar effects at 0.001 to 0.016 µg /L and 0.0003 to 0.010 µg /L,
respectively. Second generation NOEC were seen at concentrations as low as 0.0002 to 0.0003
µg/L (Lange R., 2001; Wenzel et al., 2001). Secondary endpoints including secondary sex
characteristics, testis-ova, gonadal histopathology, and VTG had NOEC in the 0.001 to 0.004
µg/L range. In both studies, and presumably due to the strong estrogenic potencies of these
compounds essentially complete feminization of males and/or male characteristics were
reported.
In order to differentiate the relationship of endpoint effects caused by a narcotic mode of action
and those mediated by an estrogenic mode of action, Wenzel et al . (2001) examined the ratio
of the acute LC50 of EE2 (1,700 µg /L) and the most significant reproduction NOEC
(fertilization success, NOEC 0.0003 µg /L). The acute-to-chronic ratio of the two is 5.73
million. This very high ratio is indicative of the mode of action associated with the synthetic
estrogen EE2 and very different from the ratios that are calculated with 4-NP (22 to 116) as is
discussed below.
In a whole lake study, Kidd (2007) and colleagues dosed a whole lake (0.005 to 0.006 µg/L)
with EE2 for 3 years. Fathead minnow populations became completely feminized, reproduction
halted, and the population collapsed. The abundance of pearl dace fish was also decreased.
These two species of fish are predominant in the lake. As noted in laboratory studies, severe
effects on gonadal tissues were found. Male gonadal tissue development was stunted or halted,
edema was noted in ovarian tissues and lesions were found in kidney tissue.
In summary, the effects on aquatic organisms by potent estrogenic substances such as E2 and
EE2 collectively add up to a predictable, cohesive signature of effects, which are consistent
with the known mechanism of action for these compounds. Natural development of male
gonadal tissue is halted, differentiation into male characteristics is blocked, male secondary sex
characteristics are prevented from developing, and gonadal tissues are severely compromised
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and damaged, reproduction within natural populations can be halted, and effects that are
mediated by estrogenic activity occur at concentrations that are millions of times lower than
those where narcotic effects occur. As a scientifically rigorous review of relevant studies shows
(see Section 2.3 below), the effects of 4-NP on aquatic organisms are very different from the
effects from potent estrogens.
2.3 4-NP affects acute lethality and chronic reproduction, growth and
development endpoints very differently than natural (E2) or synthetic (EE2)
estrogens.
The Annex XV Report for 4-NP attempts to link effects reported for 4-NP from all the studies
with fish to an endocrine disruption mode of action. Comparison of studies examining 4-NP
and natural (E2) and synthetic (EE2) estrogens demonstrates that there are great differences
between the effects that they cause to aquatic organisms.
Estrogenic activity within intact aquatic organisms due to exposure to the natural and synthetic
estrogens E2 and EE2 occurs at concentrations that are orders of magnitude below the
threshold for systemic toxicity. These exposures result in a suite of responses that are
collectively linked to the estrogenic mode of action. In a life cycle study with EE2 and
zebrafish, Wenzel et al. (2001) reported the 96-h LC50 (narcosis-based toxicity) for EE2 to
be 1700 µg/L and the NOEC for fertilization success to be 0.0003 µg/L. The authors
calculated the ratio of the two values to be 5.73 x 106. It is clear for EE2 that the
reproduction, growth and development endpoints are affected by a different mechanism of
action (endocrine) than the acute lethality (narcosis) endpoints. Given the potency of natural
and synthetic estrogens, this is expected, and hence the very high ratio is understandable.
4-NP does not affect the same endpoints in aquatic organisms in the same way that natural
and synthetic estrogens act. Data from three studies with 4-NP demonstrate that the
reproduction, growth and development endpoints are mainly indicative of narcosis-type
mode of action, possibly coupled with a very weakly estrogenic mode of action.
Using rainbow trout, Brooke et al. (1993a) reported a 96-h LC50 of 221 µg/L. Ackermann et al.
(2002) reported a no observed effect concentration (NOEC) from a one year exposure to
embryonic, larval, and juvenile rainbow trout of 10.17 µg/L based on hatching success and sex
ratio. The ratio of the lethal effect concentration (221 µg/L) to the reproductive NOEC of 10.17
µg/L is 22. Using fathead minnows, Brooke et al. (1993a) reported a 96-h LC50 of 128 µg/L.
Giesy et al. (2000) reported a NOEC of 1.6 µg/L based on egg production with fathead
minnows. The ratio of the acute lethal concentration of 128 µg/L and the NOEC of 1.6 µg/L for
the reproductive endpoint is 80. Using Japanese medaka, Kashiwada et al. (2002) reported a 96-
h LC50 of 950 µg/L. Yokota et al. (2001) conducted a 1.5 generation test with Japanese medaka
reporting a lowest NOEC for F1 sex ratio of 8.2 µg/L. The ratio of the acute lethal effect
concentration of 950 µg/L and the NOEC of 8.2 µg/L is 116. Thus, the ratio of acute lethality
LC50 values and the NOEC of the lowest apical NOEC for reproduction, growth or
development endpoints are fairly consistent across the species rainbow trout, fathead minnows,
and Japanese medaka. Across the three species, the ratios range from 22 to 116, which are
approximately 50,000- to 260,000-fold lower than the ratio of 5.73 million calculated for EE2.
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Together these data show that 4-NP has a very different mode of action than the natural and
synthetic estrogens with effects on survival resulting from narcosis being at least as critical to
the organisms exposed to 4-NP as other modes of toxicity. While biomarkers such as VTG
indicate that estrogen receptor binding is occurring while exposed to 4-NP, the endpoints
related to growth, development, and reproduction are only a factor of 22 to 116 from short term
acute toxicity and in the same range as caused by longer term lethality. Hence it is possible that
effects on growth, development, and reproduction for 4-NP are more indicative of narcosis-type
mode of action, possibly coupled with a very weakly estrogenic mode of action. Taken
together, the acute and chronic datasets for 4-NP and EE2 are very different. This conclusion is
supported by the assertion by Schwaiger et al. (2000) based on laboratory exposures of
Common carp (Cyprinus carpio) to 4-NP, “…4-NP-induced general toxic effects might
outbalance the relatively weak estrogenic effects of this compound…” as ecologically relevant
endpoints such as reproduction may be “…also due to toxic effects leading to an impairment of
the general health condition of the fish.”
2.4 4-NP operates under multiple modes of action, not just a weak estrogenic mode of
action.
While 4-NP has weak estrogenic activity that is 1000 to 1,000,000 fold less potent than 17ß-
estradiol (E2) and 17α-ethynylestradiol (EE2) (Coady et al., 2010), adverse apical effects caused
by 4-NP are not "clearly endocrine mediated”. The assumption that effects on biologically
complex, apical endpoints, such as reproduction, are mediated by a single mode of action, in this
case one that is endocrine mediated via the estrogen receptor, is not reasonable for alkylphenols.
In estrogenic mixture studies with estrogens and alkylphenols, the phenomenon of decreased fish
reproduction due to alkylphenol exposure alone was clearly not solely attributed to estrogen-like
activity (Brian et al., 2007). Investigations using gene array technologies to specifically compare
4-NP and estradiol (E2) gene transcription profiles have established that 4-NP has additional
modes of action that are independent of the estrogen receptor (Larkin et al., 2002; Ruggeri et al.,
2008; Watanabe et al., 2004). Molecular evidence in both mammalian and fish models have
demonstrated that the alkylphenols influence a greater suite of genes than estrogens (Ruggeri et
al., 2008; Watanabe et al., 2004). For example, 425 genes were differentially expressed in liver
tissue from zebrafish exposed to 10-7
M 4-NP, while 153 genes were differentially expressed in
liver tissue from zebrafish exposed to 10-7
M E2 (Ruggeri et al., 2008). Of the 30 most
differentiated genes affected by 4-NP compared to controls, only 1/3 of these genes were also
altered among E2-exposed fish, and then not all in the same direction of change (Ruggeri et al.,
2008). In mice, nonylphenol activated more genes than E2 in liver tissue, and the activated genes
in the livers of 4-NP-exposed mice were distinct from estrogen-responsive genes (Watanabe et
al., 2004). These molecular studies of gene activation illustrate that 4-NP has multiple modes of
action, of which weak estrogenic activity is one, and the apical effects noted on aquatic
organisms, such as decreased growth, survival and reproduction, are not necessarily directly or
solely attributed to 4-NP’s weak estrogen receptor binding activity.
3.0 The concern expressed in the Annex XV Report on 4-NP is for reproductive or
developmental effects in aquatic organisms caused by an endocrine mechanism;
however it does not provide adequate rationale for the leap between findings of
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weak 4-NP estrogenic activity and reproductive effects seen in some fish studies.
The endpoints that are considered to indicate an estrogen agonist mode of action are
summarized on page 49 (Table 18) of the Report. Several of these endpoints are more apical
in nature and can be observed naturally and also under various stressful conditions for fish.
Oocyte atresia normally occurs in fish as a part of the natural reproductive cycle, so an
understanding of the normal background level of this phenomenon is important. Oocyte atresia
and testicular degeneration also occur in fish when exposed not only to an estrogenic
compound, but also in response to compounds with known androgenic or anti-androgenic
modes of action (USEPA, 2007). Therefore it appears that oocyte atresia and testicular
degeneration, in and of themselves, are not indicative of a particular endocrine mode of action
(i.e. estrogen receptor agonism). It is also established that stressful situations resulting in
sustained increases in a fish’s cortisol levels can depress sex steroid concentrations, which in turn
can cause delayed gonadal development, lowered VTG levels, and depressed secondary sex
characteristics in fish (Milla et al., 2009; Aluru and Vijayan, 2009). Therefore, these
endpoints are not necessarily specific to an endocrine mediated mode of action, but will
respond when fish are stressed, whatever the cause may be.
In various studies with medaka fish, increased VTG was observed in males at concentrations
of 4-NP ≥ 5.4 µg/L, altered histopathology of the gonads (specifically occurrence of testis-
ova) generally occurred at concentrations of 4-NP ≥ 8.2 µg/L. The occurrence of testis-ova in
the gonads was more marked when fish were exposed during the time period of sexual
differentiation, which occurs before hatch. Incidences of ova-testis were less frequent and
occurred at higher concentrations when exposure to 4-NP was initiated post-hatch. Sex ratios
in medaka fish were altered toward a greater proportion of females at concentrations of 4-NP
≥ 17.7 µg/L. These results indicate that effects noted in medaka may be due to the weak
estrogen binding activity of 4-NP. However, the apical effects that have a relevance to
medaka populations (i.e. altered sex ratio), occur at concentrations well above the current
Environmental Quality Standard (EQS) of 0.3 µg/L for 4-NP in surface waters under the
Water Framework Directive as well as above the Predicted No Effect Concentration (PNEC)
of 0.6 µg/L that is listed in the REACH Dossier for 4-NP and was based on a species
sensitivity distribution model (ECHA, 2012).
In studies with fathead minnows of various ages, increased VTG was observed at
concentrations of 4-NP ≥ 8.1 µg/L, altered histopathology of the gonads and altered
secondary sex characteristics generally occurred at concentrations of 4-NP ≥ 1.6 µg/L.
Apical endpoints, such as reproduction and survival, were affected among fathead minnows,
including early life stages, at concentrations of 4-NP ≥ 14 µg/L. While behavioral effects of
4-NP exposure on male fathead minnows were described by Schoenfuss et al., 2008,
namely nest holding behavior, the magnitude of the noted effects among 4-NP-exposed fish
was low (i.e. 5-10% of 4-NP-exposed fish were outcompeted by control males) and
arguably not of biological relevance. The biological relevance of such a small effect is
questionable, especially for an endpoint as variable as nest holding behavior, which has not
been validated in a standardized fish testing guideline. Also notable in the Schoenfuss et
al., 2008 study is the fact that patchy and high mortality rates were observed among tested
fish at concentrations of 4-NP that are lower than those known to result in fish mortality
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(i.e. 58% survival in 0.15 µg/L 4-NP treatment group), thus questioning the husbandry
practices for the fish in this study. Again, the apical effects that have a relevance to fathead
minnow populations (i.e. reproduction and survival), occur at concentrations well above the
current environmental quality standard of 0.3 µg/L for 4-NP in surface waters under the
Water Framework Directive.
A quote from the section on fathead minnow results in the Annex XV Report for 4-NP (page
64) states, “Although apical effects started at similar or even lower concentrations compared
to biomarker responses, it seems very likely that they are estrogen mediated.” There really is
no support for this statement, since 4-NP has multiple modes of action apart from weak
estrogen receptor binding, as has been demonstrated by several gene array and other
molecular based mode of action studies with 4-NP (see section 2.4). There is no definitive
link with the weak estrogenic effects noted among fathead minnows (i.e. effects on VTG and
reproductive tissues) and the effects on apical endpoints (i.e. reproduction and survival)
which occur at overlapping concentrations. The most sensitive apical effect for fathead
minnows that was statistically different from controls in a validated study was based on
mortality in an ELS study (Ward and Boeri, 1991).1 There is no clear link in this study with
fathead minnow mortality and an estrogenic mode of action.
A total of four 4-NP exposure studies with zebrafish were examined in the report, however, 3
of the 4 studies were deemed a Klimisch score 3, which indicates that these studies are “not
reliable” (Klimisch, 1997). However, the endpoints from these studies are still used in a
weight of evidence evaluation of the effects of 4-NP on zebrafish. With inappropriate
controls included in the test design and high control mortality rates, the study endpoints are
compromised and should not be included in a weight of evidence approach for determining
the effects of 4-NP on zebrafish. From the one study determined acceptable for consideration
(Lin and Janz, 2006), VTG was increased among male and female fish at 100 µg/L 4-NP,
testicular tissue was skewed to earlier stages at 10 µg/L 4-NP, and the apical effects of 1) no
detectable males based on gonadal histology and 2) decreased swim up success were noted at
100 µg/L 4-NP. Altered sex ratios compared to controls were also observed at 10 µg/L 4-NP
(58% females at 10 µg/L 4-NP vs 30% females in controls), however since the sex ratio was
skewed in the control group, the meaningfulness of this observation is questionable. The
apical effects that have a relevance to zebrafish populations (i.e. altered sex ratio and
survival), occur at concentrations well above the current environmental quality standard of
0.3 µg/L for 4-NP in surface waters under the Water Framework Directive.
Various studies with rainbow trout exposed to 4-NP were summarized, and effects on
increased VTG levels were noted at concentrations as low at 1.05 µg/L. Effects on gonadal
tissues were noted at concentrations of 4-NP ≥ 30 µg/L, and effects on apical endpoints
including growth and survival, were noted at concentrations of 4-NP ≥ 10.3 µg/L.
Additionally, the effect of reduced sperm volume was noted in one study (Lahnsteiner et al.,
2005) starting at concentrations of 4-NP ≥ 0.13 µg/L; however, reproduction was not assessed
in this study, so it is unknown if this decrease in sperm volume affected reproduction levels in
1 The Ward and Boeri 1991 study with fathead minnows is scored in the Annex XV Report for 4-NP as Klimisch
score 4 (insufficient information). The Klimisch score for this study should be1. A copy of the unpublished study is
provided as Attachment B to these comments for verification.
12
rainbow trout. In addition, this endpoint is not well-studied, therefore the usefulness of the
endpoint for predicting effects for population health is not known. The lowest apical
endpoint noted in among the toxicity studies with rainbow trout was the reduced growth of
rainbow trout at 10.3 µg/L in an early life stage test (Brooke, 1993).2 This apical endpoint is
not clearly related to an estrogenic mode of action. Fish growth can be affected by various
modes of action and reductions or delays in growth are not necessarily estrogen-mediated. In
fact, the OECD Guidance Document on the Assessment of Chemicals for Endocrine
Disruption states, “… an effect solely on growth or survival, while potentially of concern
from the viewpoint of environmental risk assessment, would not lead to a conclusion that the
chemical is an ED in fish.” (OECD, 2011). Again, the apical effects that have a relevance to
rainbow trout populations (i.e. growth and survival), occur at concentrations well above the
current environmental quality standard of 0.3 µg/L for 4-NP in surface waters under the
Water Framework Directive.
In studies with guppies and mosquito fish, 4-NP exposure resulted in similar findings with
increased VTG at concentrations of 4-NP ≥ 10 µg/L, changes in gonadal tissues at
concentrations ≥ 0.5 µg/L, and effects on apical endpoints at concentrations of 4-NP ≥ 5
µg/L. While skewed sex ratios were noted at concentrations of 4-NP ≥ 50 µg/L, significant
effects on growth were noted at lower concentrations of 5 µg/L. As stated previously,
decreased growth is not specific to an estrogenic mode of action. As seen with the other fish
species, apical effects that have a relevance to viviparious fish populations (i.e. growth),
occur at concentrations well above the current environmental quality standard of 0.3 µg/L for
4-NP in surface waters under the Water Framework Directive.
Taken together, the fish studies do indicate that 4-NP has a mode of action that is consistent
with a relatively weak binding affinity to the estrogen receptor when compared to the
endogenous ligand, E2. Increased levels of VTG, alterations in gonadal histopathology, and
altered sex ratios in fish are consistent with this mode of action. However, the most sensitive
apical endpoints among fish toxicity studies with 4-NP are based on decreased growth and
survival, particularly of early life stage fishes. These endpoints are not clearly linked to an
endocrine mode of action, and are not estrogen agonist specific. There are other modes of
action associated with 4-NP as is evidenced by the suite of genes that are differentially
activated in gene array and other molecular based investigations of the mode of action of 4-
NP.
4.0 The Annex XV Report for 4-NP does not provide adequate scientific support for its
statement that amphibian data provided in the Report support a conclusion that 4-
NP causes endocrine mediated effects in amphibians and other taxa.
The Annex XV Report for 4-NP states “Results for amphibians provide indications that effects in
other taxa may be endocrine mediated i.e. caused by an estrogen-like mode of action, too.”
2 The Brooke, 1993 rainbow trout study is scored as aKlimisch score of 4 in the Annex XV Report for 4-NP due to
inaccessibility of the report. This is a US EPA study that should be scored as Klimisch 1. A copy of the report is
provided as Attachment C to these comments.
13
However, most of the amphibian toxicity studies summarized in the Annex XV Report for 4-NP
were not designed to specifically assess endocrine activity of 4-NP. Measured endpoints in the
amphibian toxicity studies commonly included tadpole growth (length and weight
measurements), metamorphic progress (an apical endpoint, sensitive to multiple types of
stressors), survival, and incidence of developmental abnormalities. None of these endpoints are
specifically indicative of estrogen receptor agonism. The report indicates that pigmentation
(melanocyte differentiation) can be estrogen responsive. However, changes in pigmentation
among both amphibians and fish can occur for multiple reasons, including as a response to stress
or altered temperature (Green and Baker, 1991; Fernandez and Bagnara, 1991); therefore
changes in pigmentation should not be interpreted as estrogen agonist specific. Taken together,
the concentrations of 4-NP resulting in effects to amphibian populations are similar to the
concentrations that cause effects to fish populations. Apical endpoints in amphibians (such as
growth, malformations, and survival) were altered at concentrations of 4-NP ≥ 100 µg/L, which
is much higher than the current EQS of 0.3 for 4-NP under the Water Framework Directive.
More specific endpoints related to estrogen receptor agonsim (i.e. gonad histology and sex ratio)
were measured in an toxicity study with 4-NP using wood frogs (Rana sylvatica) and northern
leopard frogs (Rana pipiens) (Mackenzie et al., 2003). However, there are multiple issues with
this study that indicate the data should be used with care, and these are: 1) In the study design,
wood frog exposures were not replicated due to limited availability of the eggs, 2) only two
concentrations of 4-NP were evaluated in the studies, limiting the interpretation of the data as a
full dose-response evaluation for 4-NP was not possible with the data set, 3) measured
concentrations of 4- NP indicate that the static renewal test design system was not adequate to
maintain exposure concentrations (only 9% of nominal 4-NP concentrations were measured, on
average), 4) mortality in the leopard frog study was high across all treatment groups including
controls (ranging from 40 to 58%), and 5) the authors themselves point out several uncertainties
with the intersex and sex ratio endpoints observed in the study indicating the occurrence of
intersex among controls and the unanticipated sex ratio responses of some of the chemical
exposures. Therefore, this study should not be viewed as providing definitive data for the effects
of 4-NP on amphibians in regard to estrogen specific modes of action. Additional information
for 4-NP in amphibians can be gleaned from a more recent, well-conducted, chronic study with a
close 4-NP, analogue 4-tert-octylphenol (OP). In a study with OP and Xenopus tropicalis
tadpoles, animals were continuously exposed to 0, 1.1, 3.3, 11, and 36 µg/L OP from
Nieuwkoop and Faber stage 46 to adulthood (a total of 31 weeks) with a sampling time-point at
the completion of metamorphosis as well (Porter et al., 2011). In this study, no significant
deviation from the control sex ratio was observed for either sampling period, suggesting minimal
to no effect of OP exposure on gonad differentiation. No effects in the adult frogs were observed
for mortality, body mass and size, liver somatic index, estradiol and testosterone serum levels,
sperm counts, or oocyte counts. The development and growth of oviducts, was observed in males
exposed to OP, and VTG levels were increased among juvenile male frogs (sampled at the time
of metamorphic completion), but not in adult male frogs at the close of the exposure at 31 weeks.
The transient increase in male vitellogenin and the presence of oviducts in some male frogs
exposed to OP indicates weak estrogenic activity of OP, however Porter et al (2011) concludes
14
that the highest OP concentration used, 36 µg/L, was at or below the no observed effect level
(NOEL) for toxicity in X. tropicalis.
In the Annex XV Report for 4-NP, several studies were included in the weight of evidence
evaluation of 4-NP effects on amphibians, even though these studies are not reliable for use. In
particular, the study by Kloas et al. (1999) is considered not valid as the authors failed to control
temperature during the test, had no analytical confirmation of test concentrations, had skewed
sex ratio in controls, and used improper statistics. The study with the same species by Van Wyk
et al. (2003) is considered not relevant as the authors employed intraperitoneal injection of 4-NP
as the means of dosing, which is not a relevant route of exposure for species and test material.
Therefore, these studies should not be included in the weight of evidence evaluation for 4-NP,
not should they be used to justify estrogenic effects of 4-NP in other taxa.
5.0 The scientific evidence presented in the Annex XV Report for 4-NP is in some
cases based on invalid studies and flawed study interpretation and the overall
assessment does not adequately support a case that this compound rises to a
level of concern that is equivalent to a CMR Cat. 1A, 1B, PBT or vPvB
compound.
The bulk of the Annex XV Report is devoted to presenting a case that 4-NP is a substance of
“equivalent concern”; however the scientific analysis used to make that case is circumstantial,
simplistic, incomplete, and in some cases wrong. Specific examples of these problems are
given in the discussion of the individual studies below. A full analysis of the available aquatic
toxicity data for this compound does not give rise to a conclusion that serious aquatic effects
are probable or would be comparable to a Cat. 1 or 2 CMR, PBT or vPvB substance.
In Chapter 5 of the Annex XV Report for 4-NP, the aquatic hazard assessment is presented.
The assessment introduces a number of studies mostly with fish, amphibians, and
invertebrates and identifies potential findings that might be indicative of endocrine activity.
Prior to analysis of any given study, the authors of the report gave each study a study
quality score according to the principals of Klimisch (1997). The scoring of the various
studies by the authors of the report was inappropriate in many cases. Types of errors include
scoring studies as valid (scores 1 or 2) that used no statistics, lacked replication, or had poor
control performance. Additionally, several studies were scored as “not valid” (3), but the
studies were used anyway in the Report. Studies properly scored as “not valid” (3) cannot
be used in a hazard assessment.
Some studies were inappropriately scored as “use with care” and should have been scored as
“3” or “not valid”. Ashfield et al. (1998) conducted a 400+ day study with female rainbow trout
measuring length and weight, plus ovarian weight. Test concentrations were never measured
over the course of the test, so the authors had no way of knowing test concentrations, if
concentrations were stable or drifted well beyond nominal, and did not even measure stock
solution concentrations to verify that test chemical was delivered to the test vessels. In addition,
length and weight measurements varied significantly during the 431 day test. At some time
15
points, lengths and weights were higher than controls, at other times they were lower. There
was no consistent dose related pattern of changes and most changes were within 10% of control.
The dosing is suspect and the growth effects were small and not consistent. The study should
properly be scored as “3”, not valid.
Lahnsteiner et al. (2005) exposed adult rainbow trout, and subsequently eggs, to 4-NP in the
reported range of 0.13 to 0.75 µg/L. Concentrations in test vessels and stock solutions were
never measured. The authors had no confirmation at all of the concentrations to which the fish
were exposed. Concentrations were only estimated. Effects on semen volume and fertilized egg
survival were reported at all concentrations. These data are dramatically different from all other
data with rainbow trout. For instance, Ackermann et al. (2002b) reported no effects on gonad
maturity stage, hatching success, mortality, or growth at 10.17 µg/L. Brooke (1993) reported
effects on growth at 10.3 µg/L, effects on survival at 23.1 µg/L, and no effects on time to hatch
or survival of hatched embryos at 23.1 µg/L. Jobling et al. (1996) reported effects on testis
growth at 54.3, but not 20.3 µg/L. The results of Lahnsteiner et al. (2005) should be scored as
“3” and considered not valid because the authors had no idea what concentrations were in the
test vessels and their results are not supported by other studies cited in the Annex XV Report.
The study by Ward et al. (2006) examined social behavior of juvenile rainbow trout exposed to
4-NP at 40 and 80 µg/L. Some effects on social behavior were reported at both concentrations.
However, the study employed no replication. It is not possible to demonstrate statistical
significance without replicates and thus, the study by Ward et al. (2006) should be scored as
“3”, not valid. Schoenfuss et al. (2008) conducted competitive spawning assays using fathead
minnows of either 8 months of age (Exp. 1) or 9 months of age (Exp.2). The authors reported
significant effects on competition at concentrations as low as 0.15 µg/L in Exp. 1 using 8 month
old fish, but only at 11 µg/L and higher in Exp. 2. The extraordinary difference in results
between the two nearly identical experiments that used adult fathead minnow of either 8 or 9
months of age suggests significant experimental flaws or that such competitive spawning assays
are too highly subjective to be useful for hazard or risk assessment. Due to the great differences
in results from two nearly identical experiments, the study should properly be scored as “3”, not
valid.
Arslan and colleagues (Arslan and Parlak 2007; Arslan et al. 2007) reported the results of
studies with two species of sea urchin, Paracentrotus lividus and Arbacia ixula.Spermiotoxicity,
genotoxicity, and embryo toxicity were assessed in both species of sea urchins exposed to seven
treatments of 4-NP ranging from 0.937 to 18.74 μg/l (nominal concentrations, 4-NP was not
measured). In the spermiotoxicity test, sea urchin sperm were exposed to 4-NP delivered in
dimethylsulfoxide for 30 minutes, then mixed with viable eggs. Fertilization success was
measured and the embryos were assessed for malformities and qualitatively scored in terms of
development stages. The genotoxicity test was conducted by assessing embryos for mitotic
abnormalities after 5 hours of exposure to 4-NP at the concentrations listed above. In the
embryo toxicity test, sperm and eggs were added to the treatments together and exposed to 4-NP
treatments throughout 72 hours of embryonic development. Spermiotoxicity effects on larval
development of P. lividus were observed at all concentrations of4-NP down to 0.937 μg/l
(Arslan et al. 2007). These results are in contrast to other work reporting effects on the sea
urchin, P. lividus, in which sea urchin sperm was exposed to 4-NP for 72 hours after which an
16
EC50 of 270 μg/l based on sperm toxicity was recorded (Ghirardini et al. 2001). In light of
conflicting data, there is uncertainty associated with the results with sea urchins reported by
Arslan et al. (2007). Embryotoxic effects among both species of sea urchins were observed at
all concentrations of 4-NP down to 0.937 μg/l (Arslan et al. 2007; Arslan and Parlak 2007).
However, due to potential solvent effects and a lack of test chemical measurements in the static
tests (in which test chemical concentrations are likely to be less consistent than in flow-through
tests), the use of the Arslan et al. (2007) and Arslan and Parlak (2007) studies for hazard or risk
assessment should be considered with caution.
6.0 As stated under Article 57(f) scientific evidence of probable serious effects to human
health or the environment is necessary to classify a compound as a SVHC; the very
low estrogenic potency, multiple modes of action, and low environmental
occurrence and concentrations do not support a conclusion of 4-NP of probable and
serious environmental effects from this compound.
Categorization as SVHC should be reserved for substances for which the weight-of-evidence
supports a finding of probable and serious effects. As discussed previously, 4-NP, at most, has
weakly estrogenic activity and the available data do not support a conclusion that this compound
is of equivalent concern to a CMR Cat. 1 or 2, PBT or vPvB compound. As discussed below,
concentrations of 4-NP in European waters are significantly below NOEC and LOEC values for
adverse apical effects found in studies on 4-NP in aquatic species. In addition, studies that
examine estrogenically active substances in wastewater treatment plant effluent found that if
detected, 4-NP and alkylphenols generally, contributed only minimally to the aggregate
estrogenicity.
6.1 Concentrations of 4-NP in European surface waters do not support concern for
probable and serious effects in aquatic species
4-NP Monitoring data were collected between 2007 and 2009 in the context of the water
framework directive (DGEnv, 2009). In the water dissolved fraction mean concentrations of
0.040 µg/L (maximum 0.460 µg/L, median 0.030 µg/L) were observed. These dissolved
concentrations are well below all NOEC and LOEC from the screening, short term reproductive,
and life cycle studies with 4-NP that ranged from approximately 1 to 180 µg/L. These data
provide further support to the evidence that 4-NP does not cause “probable serious effects” to the
environment that are equivalent to CMR Cat. 1 and 2, PBT, or vPvB substances.
6.2 Contribution of 4-NP to Measured Estrogenicity of Effluents or Surface Waters
Numerous studies have examined the extent to which measurable estrogenic activity in
wastewater treatment plant (WWTP) effluents is attributable to natural and synthetic compounds
(e.g., E2 and EE2) or to other constituents (e.g., 4-NP or other alkylphenols).
Desbrow et al. (1998) used a sample fractionation system to isolate fractions of WWTP effluent
samples that were tested for estrogenicity using a YES assay. Estrogenic fractions were found to
contain natural and synthetic hormones including E2, estrone, and EE2. Alkylphenolic
compounds were not detectable in the estrogenic fractions.
17
Most studies employed in vitro screening assays such as YES assays and concentration
measurements from whole effluents to assess relative contributions to estrogenicity. Korner et al.
(1999) and Aerni et al. (2004) reported that the contribution of 4-NP in German WWTP effluents
was <3% and 2%, respectively, of the total estrogenicity. In WWTP effluents in the US, Drewes
et al. (2005) reported that nearly all of the estrogenic activity was attributable to natural and
synthetic hormones. Similar findings have been reported for WWTP effluents in Australia, New
Zealand, and China, (Leusch et al., 2006; Jin et al., 2007; Lu et al., 2011).
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